In recent years there has been an increased interest in the use of microorganisms for environmental restoration. The usefulness of microorganisms with a diversity of metabolic activities in wide ranging applications coupled with advances in the technology have led to successful demonstrations in the ever-expanding bioremediation field. Bioremediation is the degradation or stabilization of contaminants by microorganisms (e.g., bacteria, fungi, actinomycetes, and cyanobacteria). It is one of the oldest in-situ remedial processes, having first been practiced commercially in 1972. Since that time the technology has matured from a novel process to an important and recognized, economical method of remediating soils and groundwater. The targets for bioremediation continue to expand to the higher weight petroleum compounds (greater than C-15), and chlorinated, nitrated, oxygenated and
saturated organic compounds.
The application of environmental biotechnology as a successful remediation tool depends on the ability to stimulate or enhance specific activity of indigenous or introduced microorganisms. The challenge has been to enhance the activity of these microorganisms and develop means to bring the contaminant into direct contact with the organisms to achieve optimal bioremediation.
During the bioremediation process, under ideal conditions, organic compounds are converted by soil microorganisms to carbon dioxide, water, biomass and inorganic salts (e.g. chlorinated compounds). Numerous factors affect both the rate and the extent of petroleum derived hydrocarbon mineralization in contaminated soils and water. These factors are:
- nutrient level
- contaminant and soil characteristics
Biodegradation is a safe, effective, and economic alternative to traditional methods of remediation. Bioremediation is now often used in conjunction with a wide range of physical and chemical technologies to provide an optimal remediation strategy.
Bioremediation alternatives are currently being researched using a combination of physical, microbiological, chemical and molecular based methods. Intrinsic bioremediation, the remediation of contaminants by naturally-occurring biological processes at background rates, is being studied as an avenue for the restoration of contaminated media in cases where other methods of remediation are not
Bioremediation research in the laboratory includes isolating and characterizing naturally-occurring microbial isolates with bioremediation potential, developing genetically engineered microorganisms and enzymes with contaminant degrading capabilities, studying their catabolic activity in contaminated media through bench-scale experiments, and monitoring the progress of bioremediation through chemical
analysis and toxicity testing in chemically-contaminated media. Field applications of bioremediation include the in-situ stimulation of microbial activity by the addition of microorganisms and nutrients and the optimization of environmental factors and the ex situ restoration of contaminated media by landfarming and composting methods.
Impact of In Situ Chemical Oxidation on Native
Biological Populations — Review of Case Studies
James Studer (ChemRem International LLC, Albuquerque, New Mexico, USA)
Greg Davis (email@example.com) and Brett Baldwin
(Microbial Insights, Rockford, Tennessee, USA)
Gary Cronk (JAG Consulting Group, Santa Ana, California, USA)
Soil, bedrock, and groundwater environmental treatment technologies have evolved over the previous decade to include use of relatively aggressive techniques involving chemicals that can dramatically alter, sometimes permanently, the physical and geochemical nature of the subsurface materials targeted for treatment. Subsurface microorganisms can also be adversely impacted by the chemical directly or indirectly as a result of physical and geochemical changes in the microbial environment. Of current interest to the environmental remediation industry is the improvement in understanding of how one class of aggressive technologies – chemical oxidation – are impacting native microbial populations, at the basic science level and relative to the practical concerns of the remediation engineer.
Chemical oxidation technologies involving sodium persulfate (as a primary oxidant) and associated activation chemicals that can generate oxidative and reductive radicals are being increasingly applied across the globe. The native microbial response to use of these chemicals (specifically sodium persulfate, calcium peroxide, sodium hydroxide, and hydrogen peroxide) is reported by considering several recent groundwater contamination projects that have benefited from advances in molecular biological characterization technologies and protocols.
Project A is located in California and involved a complex mix of chlorinated and nonchlorinated solvents and 1, 4 – Dioxane.
Project B is situated in Mississippi and involved a mix of wood treating wastes (PCP and PAHs).
Project C involved diesel and gasoline petroleum hydrocarbons in fractured bedrock of Alberta and included a Bio-Trap Sampler baited with 13C labeled benzene. The heavy isotope labeled was used as a model to track the fate of petroleum hydrocarbons into biomass and inorganic carbon sinks.
Finally, Project D is located in California and involved a benzene-contaminated aquifer. Phospholipid fatty acid (PLFA) analyses were conducted prior to and following addition of sodium persulfate (and one or more of the following chemicals: engineered calcium peroxide, hydrogen peroxide, sodium hydroxide) were used to elucidate changes in viable biomass, microbial populations, and response to oxidative stress. Quantitative Polymerase Chain Reaction (qPCR) analyses were conducted to enumerate specific genes responsible for Contaminant biodegradation.
In general, results from these four projects indicate that microbial populations are temporarily suppressed one to two orders of magnitude and population shifts and degree of stress occur. Ultimately, native populations are robust and implications for effective engineering of coupled or sequential chemical oxidation and aerobic and/or even anaerobic bioremediation are encouraging.
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Nutrient-Enhanced Attenuation Technology (NEAT®) is eGeo’s trademarked process employed on remediation of chlorinated solvents, PERC, and related COCs. The anaerobic and aerobic process of NEAT® injects vapor-phase nutrients through a unique delivery system, which converts TCE to TCE-epoxide. TCE-epoxide is highly unstable and breaks down into chloride salt and carbon dioxide.
Using a co-metabolite,such as butane, propane or methane, with the vapor-phase nutrients, NEAT® is highly effective on chlorinated daughter products such as DCE and vinyl chloride. The co-metabolite stimulates the growth of methanotrophic bacteria present in soil and saturated sediments.
The NEAT® remediation system consists of a self-contained gas delivery system with a separate injection module tailored to the specific conditions of the site to accomplish the groundwater remediation.
NEAT® is based on PHOSter II™, an in-situ process consisting of the controlled addition of vapor-phase nutrients to dramatically increase the microorganism population. This patented process, originally developed by Department of Energy scientists at the Savannah River Technology Center in Aiken, South Carolina, is a pulsed injection system that delivers a measured and predetermined combination of phosphorous, nitrogen, oxygen and a carbon-based food source (electron donor), if required.
The State of Florida Department of Environmental Protection Innovative Technologies Program accepted our PHOSter II™ process in 2002, and it has been successfully employed on projects since then in Alabama, Florida, Georgia, Pennsylvania, and South Carolina. View Certification Letter
During the bioremediation process organic compounds are converted by soil microorganisms to carbon dioxide, water, biomass and inorganic salts (e.g. chlorinated compounds). Numerous factors affect both the rate and the extent of petroleum derived hydrocarbon mineralization in contaminated soils and water. These factors are:
> nutrient level
> contaminant and soil characteristics.
Critical to establishing and maintaining the activity of the bacteria in the site soil is the control of the inorganic nutrients required by the bacteria for the cellular metabolism. Bacteria require both nitrogen (N) and phosphorus (P) in order to carry out cellular functions. Nitrogen is required by the organism for microbial wall components, nucleic acids and proteins. Nitrogen makes up nearly 15% of the molecular composition of a bacterial cell; any limitation on the nitrogen seriously impacts the metabolic functions of the organisms. Bacteria utilize only that nitrogen which is in the ammonia (NH4) form, N in other forms
cannot be incorporated into the organisms organelles. Instead, improperly monitored, other N sources may lead to field and laboratory results which mislead the practitioner of bioremediation into believing that the inorganic N level is satisfactory.
Phosphorus, in the form of inorganic phosphates, is used by the microorganisms in the synthesis of phospholipids and nucleic acids. Phosphorus is also essential for the energy transfer reactions of ATP. The enzymes that hydrolyze the phosphate ester are present in nearly all organisms. In P limiting environments the metabolism of microorganisms decreases sharply, reducing their capacity to utilize the hydrocarbons in the soil as an organic carbon source. Inorganic phosphorus sources should be applied with the nitrogen sources in order to stimulate and maintain the remediation project.
The injection of a carbon-based food source, such as butane, methane or propane is useful when developing a substantial bacterial cell population where the contaminants are recalcitrant and not available for direct assimilation by bacteria. Propane(C3H8) is a very simple carbon source and can readily be converted to energy and cell mass by bacteria. The larger population will secrete enzymes of sufficient quantity to break-down the more complex compounds into forms available for direct assimilation by the population.
All three of these constituents are pulsed into the saturated zone using air as a carrier gas. Through pulsed injections we have been able to affect a greater radius of influence than using continuous injection. Also, pulsing injections encourages the development of more “enzyme active” sites, where potentially more contaminants can be broken down between pulses. We have also found that accurate control of each of the additives is necessary in treatment and that the mixture can be adjusted when targeting specific chemical compounds.
The gas injection volumes and rates are far below the levels that could induce air sparging. Because nutrients and food are injected in a mixture of air, the term “biosparging” (defined as “air and specific gases injected into saturated sediments”) may be applied to the PHOSterTM process. Injection points are sealed against air, gases or vaporized contaminants venting to the atmosphere.
Advantage of vapor-phase injection: Quicker and more thorough dispersal of nutrients can occur if they are injected in the vapor-phase, rather than as liquids or solids, especially at sites where permeability of the soil is low. For a petroleum test site in Aiken, South Carolina, where
significant contaminant reductions were obtained in 131 days, the soil permeability was relatively low: 10-8 cm2.
1) Each remediation unit has a programmable logic controller (PLC) to remotely and continually control each injection point’s pulsed cycle.
Because of low injector flow rates (0.5 to 2.0 cfm), our process doesn’t require any off-gas treatment or recovery wells.
2) eGeo’s systems remain on-site for the duration of the cleanup effort allowing for more effective process control.
3) Significantly lower electrical power costs vs. conventional processes (such as Dual-Phase/Liquid Ring.
4) No air emissions monitoring is necessary for the PHOSter II™ system since injection gas flow rates are not high enough to volitilize appreciable amounts of VOC.
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Bioreduction of PCE
Currently, there are 13 strains of bacteria known to reduce PCE to TCE to DCE, one other which is known to reduce PCE to TCE. One of the 13 is known to reduce PCE to ethene. All of these species reduce using direct anaerobic dehalogenation; this process is also called halorespiration. The bacteria are using the chemical as an energy source. Hydrogen is the electron donor and PCE/TCE/DCE/VC are the electron acceptors.
Of these bacteria most are considered fairly abundant. Of course, the one that can reduce from PCE to ethene and has the highest know yield rate (Dehalococcoides ethenogenes, Strain 195) is only considered common and, hence, you can’t count on it to be on your site, when needed. Most of the 14 are anaerobes. Some are obligate anaerobes and will only function under anaerobic conditions. Some are strict anaerobes and may die on contact with oxygen. One of the strains is a facultative aerobe (MS-1) but will only reduce chlorinateds under anaerobic conditions. All prefer a neutral pH.
The literature indicates that you can reduce PCE at most sites with on-site bacteria under anaerobic (DO<1 mg/L), neutral (6.5<pH<7.5), reductive conditions (ORP<-300 mV). In most cases, the process will stall at DCE and go no further (unless Strain 195 is available); the main problem is lack of control over the state of the aquifer. This is the main reason PCE is not naturally attenuating. Most aquifers that are impacted with only PCE/TCE will probably remain aerobic (unless you have organic rich soil) since the organic loading is too low to cause much of an oxygen demand. This is also one reason you get PCE/TCE reduction in mixed plumes. The non-chlorinated organics provide the oxygen demand to drive the aquifer anaerobic. If you successfully bio-reduce PCE/TCE on a site then the site must be in an anaerobic state and will stall at the DCE unless Strain 195 is available. If the aquifer can be manipulated to go aerobic, DCE and vinyl would reduce. Several prevalent aerobic species will directly assimilate DCE and VC.
Bioreduction of TCE
This section describes the great interest in methanotrophic bacteria in bioremediation. Methanotrophic bacteria have a ubiquitous distribution in the environment and the use of natural gas or methane with other nutrients to stimulate their bioremediation activities through methane monooxygenase is a remediation option. These two features allow for a relatively efficient, inexpensive, and safe means to manipulate the environment to accelerate bioremediation.
The methanotrophs are aerobic bacteria that oxidize methane as an energy source (carbon) through the enzyme methane monooxygenase (MMO). This MMO can cometabolize or transform nongrowth substrates by either growing or resting cells. Cometabolism is a result of nonspecific MMO activity towards organic compounds that do not serve as carbon or energy sources. While many cometabolizing bacterial species have been identified, the best studied are the methanotrophs. The reason for this is that methanotrophs are ubiquitous and can cometabolize many aliphatic compounds, alkanes, and aromatic compounds. Methanotrophs have been intensely studied for use in degrading chlorinated solvents, most notably trichloroethylene, to environmentally acceptable concentrations in soils, sediment, and groundwater. Stimulation of methanotrophic bacteria is accomplished through the addition of methane and other gaseous nutrients resulting in an increase in contaminant biodegradation and biotransformation. The composition of gaseous nutrients used with methane is dependent on the characteristics of the site geochemistry and microbiology.
Several species of methanotrophs have been isolated that produce methane monooxygenase and participate in the cometabolic degradation of TCE (Koh et al., 1993). However, soluble methane monooxygenase (sMMO) is present in some methanotrophic bacteria that has been found to be responsible for oxidizing a wide range of carbon substrates (Cardy et al., 1991). No sMMO has been isolated or observed in type I methanotrophs (Soh et al., 1993). Moreover it appears that the sMMO is limited to a few species of type II and type X methanotrophs. The sMMO has been found through its non-specific activity to effect oxidations, dechlorinations, condensations, and rearrangements of meta-chlorotoluene, phenol, chlorofluorobenzenes, and mono and dichlorobiphenyls via it’s non-specific activity. Soluble MMO is believed to be found only in type II and type X methanotrophs during copper limiting conditions (Murrell, 1992). Although membrane and particle associated MMO has been reported, the rates of chlorinated solvent degradation with these organisms is significantly less than sMMO bacteria (DiSpirito et al., 1992). These solvents have included halogenated aliphatic compounds such as trichloroethylene (TCE). Methanotrophs can utilize nitrogen as nitrate, ammonia, and molecular nitrogen while demonstrating TCE transformation capacity (Chu and Alvarez Cohen 1996). This article will concentrate primarily on methanotrophic bacteria applications shown to be effective in chlorinated solvent bioremediation with a focus on TCE.
Wilson and Wilson (1985) demonstrated that TCE is susceptible to cometabolism by soil communities enriched with natural gas. Fliermans et al., (1988) and others demonstrated that cultures enriched with methane and propane could cometabolically degrade a wide variety of chlorinated aliphatic hydrocarbons including ethylene, 1,2-cis-dichloroethylene (c-DCE); 1,2-trans-dichloroethylene (t-DCE); and vinyl chloride (VC); toluene; phenol and cresol. Investigations that utilized methods to stimulate microorganisms in the subsurface vadose zone and groundwater found that both aerobic and anaerobic processes contributed in removal of TCE as revealed by detection of TCE oxidation and anaerobic transformation products, c-DCE; t-DCE; and VC (Hazen et al., 1993; Fliermans et al., 1994). These two studies proved that bioremediation provides a powerful means of restoring contaminated aquifers. McCarty and Semprini (1994) concluded from laboratory and field studies that cometabolic transformation of TCE was strongly tied to methane utilization. When methane addition was stopped, TCE transformation ceased.
It is now well recognized that TCE and other chlorinated aliphatic compounds can be degraded by a diversity of bacteria including methanotrophs (Little et al., 1988), selected methanogens (Bouwer and McCarty, 1984), and species of Pseudomonas (P. cepacia, P. mendocina and P. putida) capable of also degrading aromatic compounds (Nelson et al., 1988). Ensley (1991) has demonstrated a linkage between TCE degradation and aromatic metabolism in P. cepacia G4, P. mendocina and P. putida. Easign et al., (1992) reported that pure cultures of Xanthobacter spp. cometabolized TCE with the utilization of propylene as a carbon and energy source presumably using the enzyme alkene monooxygenase. Fliermans, et al (1988) and Bowman et al., (1993) have shown that enrichments for methanotrophs in subsurface samples collected from the Savannah River Site in South Carolina stimulate the microbial degradation and complete mineralization of TCE and other chlorinated aliphatic compounds both in the laboratory and in situ. Propane utilizers or propanotrophs that also exhibit non-specific oxidase activity may also be used for bioremediation of VOCs. Where mixtures of chlorinated aliphatic hydrocarbons including 1,1,1-trichloroethane, are present propane may be the stimulant of choice using air-sparging technology (Tovanabootr & Semprini, 1998). However, methanotrophs are optimal in bioremediation when TCE is the primary contaminant of concern.
Kinetics of TCE, cDCE and VC Dechlorination by Dehalococcoides-
Containing Cultures in the Presence of Chlorinated Ethanes
Winnie Chan (firstname.lastname@example.org) and Frank Löeffler
(Georgia Institute of Technology, Atlanta, Georgia, USA)
Elizabeth Edwards (University of Toronto, Toronto, Ontario, Canada)
Chlorinated solvent-impacted sites typically contain multiple contaminants and their breakdown products. This complicates potential remediation efforts because certain contaminants may interfere with the biotransformation or others. For example, 1,1,1-trichloroethane (1,1,1-TCA) and its daughter product 1,1-dichloroethane (1,1,-DCA) are common co-contaminants with trichloroethene (TCE) at field sites and have been shown to inhibit TCE-dechlorinating mixed microbial cultures.
The main objective of this research is to develop a better understanding of how 1,1,1-TCA and 1,1-DCA inhibit dechlorination of chlorinated ethenes, and at what concentration inhibition is observed. The following questions were addressed:
(1) Are the chlorinated ethanes inhibitory to the key dechlorinating organisms, or to other organisms in the consortium?
(2) Is inhibition acting directly on the enzymes (i.e., the reductive dehalogenases) responsible for dechlorination? Laboratory-scale measurements of the rates of the critical VC to ethene dechlorination step in the presence of different Concentrations of putative inhibitor (1,1,1-TCA or 1,1-DCA) were carried out using Dehalococcoides-containing whole cell and cell-free extract assays. Results for cell-free extract reductive dechlorination assays with VC as substrate generally gave a good fit to Michaelis-Menten enzyme kinetic models. When these same assays were repeated with the addition of 1,1,1-TCA or 1,1,-DCA, apparent Vmax (maximum reaction rate) and KM (half saturation constant) and inhibitor constants (Ki) for each substrate and inhibitor combination were calculated.
The main conclusions from these experiments are that 1,1,1-TCA inhibits VC
dechlorination in several distinct Dehalococcoides-containing cultures at concentrations frequently observed at contaminated sites, and is a much stronger inhibitor than 1,1-DCA. Moreover, on the basis of cell free extract experiments, 1,1,1-TCA appears to act directly on the enzymes (reductive dehalogenases) responsible for VC dechlorination.
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EPA 600/R-08/107 | September 2008 | www.epa.gov/ada
“Natural Attenuation of the Lead Scavengers 1,2-Dibromoethane (EDB) and 1,2-Dichloroethane (1,2-DCA) at Motor Fuel Release Sites and Implications for Risk Management”
Monitoring for concentrations of EDB in ground water can be a major cost of risk management at gasoline spill sites. The MCL for EDB is one hundred fold lower than the MCLs for Benzene or 1,2-DCA. Because the MCL for EDB is so low, not all analytical methods can detect EDB when it is present at its MCL. The EPA Method that is most commonly used to analyze for gasoline constituents in ground water (Method 8260B) has a detection limit for EDB of approximately 3.0 µg/L, which is sixty fold higher than the MCL. As a result, Method 8260B cannot be used to document that ground water is free of contamination from EDB. In contrast, EPA Method 8011 has a method detection limit for EDB of approximately 0.01 µg/L, which is sufficiently sensitive to measure EDB at its MCL.
Method 8260B would have only discovered 40% of the survey sites with concentrations of EDB above its MCL. At sites where benzene is the primary risk driver, Method 8260B would be appropriate to monitor the quality of ground water during active remediation. However, to determine if the site has reached the MCL for EDB, it is necessary to use Method 8011 or its equivalent.
Analysis of EDB can be included in the routine analysis of BTEX compounds by Method 8260B at minimal extra cost. In contrast, monitoring for EDB by 8011 and monitoring for the BTEX compounds by Method 8260B can essentially double the total cost of analysis. The selection of one method over the other depends on the goals and priority in risk management. Figure 3.5 presents the practical consequence of the difference in detection limits using Method 8260B or Method 8011. It compares the distribution of EDB that was determined in the EPA/ASTSWMO survey to the distribution that would have been discovered if the survey had been conducted using Method 8260B. Method 8260B would have discovered only 40% of the sites with concentrations of EDB above the MCL.
EDB Remediation Case Study
Anaerobic Biodegradation of Ethylene Dibromide and 1,2-Dichloroethane in the Presence of Fuel Hydrocarbons
James K. Henderson*†, David L. Freedman†, Ronald W. Falta†, Tomasz Kuder‡ and John T. Wilson§
Department of Environmental Engineering and Earth Sciences, Clemson University, Clemson, South Carolina 29634-5002, University of Oklahoma School of Geology and Geophysics, Norman, Oklahoma 73019, and Robert S. Kerr Environmental Research Center, U.S. Environmental Protection Agency, Ada, Oklahoma 74821-1198
Environ. Sci. Technol., 2008, 42 (3), pp 864–870
Publication Date (Web): December 19, 2007, Copyright © 2008 American Chemical Society
* Corresponding author phone: (704) 281-8840 ; fax: (864) 656-0672; e-mail: email@example.com.
Field evidence from underground storage tank sites where leaded gasoline leaked indicates the lead scavengers 1,2-dibromoethane (ethylene dibromide, or EDB) and 1,2-dichloroethane (1,2-DCA) may be present in groundwater at levels that pose unacceptable risk. These compounds are seldom tested for at UST sites. Although dehalogenation of EDB and 1,2-DCA is well established, the effect of fuel hydrocarbons on their biodegradability under anaerobic conditions is poorly understood. Microcosms (2 L glass bottles) were prepared with soil and groundwater from a UST site in Clemson, South Carolina, using samples collected from the source (containing residual fuel) and less contaminated downgradient areas. Anaerobic biodegradation of EDB occurred in microcosms simulating natural attenuation, but was more extensive and predictable in treatments biostimulated with lactate. In the downgradient biostimulated microcosms, EDB decreased below its maximum contaminant level (MCL) (0.05 µg/L) at a first order rate of 9.4 ± 0.2 yr−1. The pathway for EDB dehalogenation proceeded mainly by dihaloelimination to ethene in the source microcosms, while sequential hydrogenolysis to bromoethane and ethane was predominant in the downgradient treatments. Biodegradation of EDB in the source microcosms was confirmed by carbon specific isotope analysis, with a δ13C enrichment factor of −5.6‰. The highest levels of EDB removal occurred in microcosms that produced the highest amounts of methane. Extensive biodegradation of benzene, ethylbenzene, toluene and ortho-xylene was also observed in the source and downgradient area microcosms. In contrast, biodegradation of 1,2-DCA proceeded at a considerably slower rate than EDB, with no response to lactate additions. The slower biodegradation rates for 1,2-DCA agree with field observations and indicate that even if EDB is removed to below its MCL, 1,2-DCA may persist.
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ITRC (Interstate Technology & Regulatory Council). 2007. Remediation Technologies for
Perchlorate Contamination in Water and Soil. PERC-2. Washington, D.C.: Interstate
Technology & Regulatory Council, Perchlorate Team. www.itrcweb.org.
Perchlorate, an anion, consists of one chlorine atom bonded to four oxygen atoms (ClO4–) and is
both naturally occurring and manmade. Highly soluble and mobile in water, perchlorate is
generally very stable in the dissolved state. Most of the attention focused on perchlorate has
concerned its presence in groundwater and surface water. However, perchlorate can also be
found in soil and vegetation and has entered the human and environmental food chains.
Perchlorate occurrence in drinking water and food supplies is a human health concern because it
can interfere with iodide uptake by the thyroid gland and result in decreased thyroid hormone
Past management practices were not concerned with the release of perchlorate to the
environment because it was not recognized or regarded as a contaminant of concern. Widespread
perchlorate presence in the United States was observed after the spring of 1997 when an
analytical method was developed with a quantitation level of 4 parts per billion. Subsequent
advances in analytical chemistry have proven perchlorate to be more widespread in the
environment than previously thought.
The success or failure of a treatment technology often depends on having a complete
understanding of the nature and extent of the release. Site investigators start with a conceptual
site model, which is gradually refined through sampling and other investigative techniques.
A variety of remediation technologies are currently commercially available and are being used
for perchlorate remediation. Most of these remediation technologies fall into two broad
categories: physical and biological treatment processes.
Physical treatment processes remove perchlorate from impacted media but do not alter its
chemical composition. Considerable progress has been made in developing innovative physical
processes for removing perchlorate from drinking water, groundwater, and surface water. Some
technologies are proven and commercially available, while others are still in the research and
Ion exchange, the most proven and widely accepted physical process technology for perchlorate
treatment, is a process by which ions of a given species are displaced from an insoluble exchange
material by ions of a different species in solution. Perchlorate selective ion exchange targets
perchlorate using conventional ion exchange resin beds with specially designed resins that
preferentially remove perchlorate anions.
Biological degradation of perchlorate involves reducing bacteria, which are widespread in the
environment. Perchlorate-reducing bacteria have the ability to grow in either the presence or
absence of air, provided proper nutrients are available in the environment. Both in situ and ex
situ biological treatment systems have been applied at full scale to treat perchlorate.
The success of in situ enhanced anaerobic bioremediation largely depends on the presence of
appropriate perchlorate-reducing bacteria and the ability to stimulate sufficient growth in situ and
activity to degrade perchlorate to the extent and rate that meets the intended remedial objectives.
The ability to create the appropriate reducing conditions or to properly distribute the electron
donor to maximize contact with the contaminant and the microbes are common issues when
applying enhanced anaerobic bioremediation. Determining the potential for complete anaerobic
biological reduction using substrate addition is perhaps the most difficult question to answer in
the site-screening process.
Initially, a site can fall
into one of three microbiological categories. For some sites, appropriate perchlorate-reducing
microorganisms are present (generally believed to be widespread in the environment [Coates et
al. 1999, Logan 2001]), geochemical conditions are appropriate for their growth, and evidence of
anaerobic biological reduction is observed. In the second type of site, appropriate perchloratereducing
microorganisms are present but at an insufficient level of activity, often due to an inadequate amount of substrate (electron donor) being present to support reduction of
perchlorate. Because many aquifers are obligotrophic and aerobic, this scenario is very common
in groundwater. A third type of site is also possible in which appropriate perchlorate-reducing
bacteria are absent. However, this situation appears to be very uncommon based on laboratory
studies and field experience (e.g., Hatziner 2005, Coates et al. 1999). In the first two cases,
biostimulation alone (i.e., addition of an appropriate substrate) can often be applied with success.
In the second case, more substrate may be required to reduce oxygen and nitrate in addition to
perchlorate, but the likelihood of success is good as long as perchlorate-reducing strains are
present. Substrate amendment and bioaugmentation would be necessary for in situ treatment of
the third type of site. However, as previously noted, this scenario is uncommon due to the
general ubiquity of perchlorate-reducing bacteria.
Enhanced in situ anaerobic bioremediation can be an effective method of degrading perchlorate
in groundwater. Advantages of enhanced anaerobic bioremediation include complete
mineralization of perchlorate in situ with little impact on infrastructure and relatively low cost
compared to more active engineered remedial systems.
Enhanced in situ anaerobic bioremediation can be implemented to provide source area or
dissolved plume treatment or containment, or a combination of source area and dissolved plume
remediation can be used. Enhanced bioremediation will be subject to the same difficulties
associated with mass transfer limitations of a continuing source and preferential flow paths in
heterogeneous formations. The single largest difference between conventional remedial
technologies and enhanced bioremediation may be that enhanced bioremediation, if properly
implemented, can maintain effectiveness over a longer period of time at a lower overall cost.
Typical system configurations and associated remedial action objectives that engineered
anaerobic bioremediation may be used to address include the following:
> Source Zone Treatment. Remediation of source zones where good substrate/contaminant
contact is possible.
> Plume Containment Using a Biologically Reactive Barrier. Reduction of mass flux from a source zone or across a specified boundary.
> Plume-wide Restoration. Total treatment of an entire dissolved plume. In some cases, several
approaches may be combined. For example, a source area may be targeted for remediation using a grid configuration, combined with a linear barrier configuration upgradient from a downgradient point of compliance.
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